Effect of zinc oxide nanoparticles in Daphnia magna

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Universidade de Aveiro 2012 Departamento de Biologia

Sílvia

Ferreira Lopes

Efeito de nanopartículas de óxido de zinco

em Daphnia magna

Effect of zinc oxide nanoparticles in

Daphnia magna

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Universidade de Aveiro 2012 Departamento de Biologia

Sílvia

Ferreira Lopes

Efeito de nanopartículas de óxido de zinco

em Daphnia magna

Effect of zinc oxide nanoparticles in

Daphnia magna

Dissertação apresentada à Universidade de Aveiro para cumprimento dos requisitos necessários à obtenção do grau de Mestre em Biologia Aplicada – Ramo de Toxicologia e Ecotoxicologia, realizada sob a orientação científica da Doutora Susana Patrícia Mendes Loureiro, Investigadora auxiliar do Departamento de Biologia e CESAM da Universidade de Aveiro.

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o júri

Presidente Professora Doutora Maria Adelaide de Pinho Almeida Professora Auxiliar, Departamento de Biologia, Universidade de Aveiro

Vogal – Arguente Professora Doutora Isabel Maria Cunha Antunes Lopes

Investigadora Auxiliar, Departamento de Biologia e CESAM, Universidade de Aveiro

Vogal – Orientador Doutora Susana Patrícia Mendes Loureiro

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agradecimentos _{Aos meus pais, por todos os esforços feitos, pois sem eles}

nada disto seria possível.

Ao Vincent, por todo o apoio, compreensão e paciência que teve para comigo durante os altos e baixos desta etapa. Merci pour tout ♥.

À minha orientadora, Susana Loureiro, por toda a disponibilidade, ajuda e conselhos. Agradeço também à Fabianne pela ajuda que sempre me disponibilizou.

A todos os meus colegas de laboratório principalmente ao Carlos, Sofia, Rita, Cátia e Ariana que sempre me apoiaram, ajudaram e fizeram com que todo o tempo passado no laboratório fosse um tempo muito bem passado.

E finalmente, às “minhas” dáfnias por amavelmente terem “colaborado” para a realização deste trabalho.

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palavras-chave _{Nanotecnologia, nanoparticulas de óxido de zinco, Daphnia} magna

resumo O rápido desenvolvimento da nanotecnologia com o consequente

aumento na produção de nanopartículas e nanoprodutos oferece muitas oportunidades mas também muitos desafios. A nanotecnologia tem vindo a ser descrita como uma área multidisciplinar que visa desenvolver uma variedade de nanoparticulas para aplicações medicinais e industriais.

As propriedades que trazem às nanoparticulas especial atenção – pequeno tamanho, elevada área de superfície e consequente elevado grau de reatividade – podem também torná-las potencialmente perigosas para a saúde humana e para o ecossistema. A avaliação dos potenciais riscos inerentes à exposição das nanoparticulas torna-se portanto uma investigação de prioridade antes que estas sejam aplicadas em produtos comerciais e libertadas para o ambiente.

Os ambientes aquáticos (de água doce e marinho) são considerados como potenciais destinos das nanoparticulas libertadas para o ambiente através de fontes diretas e/ou indiretas, expondo assim os organismos aquáticos a elevados níveis de contaminação.

As nanoparticulas de óxido de zinco (ZnO-NPs) são uma das nanoparticulas mais utilizadas numa vasta gama de produtos comerciais (ex: protetores solares, cosméticos e tintas) e a sua produção estima-se que irá continuar a aumentar nos próximos anos. Em consequência, o risco de contaminação aquática por parte destas nanoparticulas irá forçosamente aumentar.

Estudos toxicológicos já demonstraram que as ZnO-NPs exercem efeitos tóxicos em vários organismos, como por exemplo, em crustáceos, algas e bactérias. Os efeitos tóxicos das nanoparticulas são complexos e podem estar dependentes de vários fatores, tais como, o organismo-teste, fatores abióticos (pH, salinidade, dureza da água e presença de matéria orgânica), propriedades físico-quimicas das nanoparticulas, processos de adsorção, presença de outros contaminantes, entre outros.

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Os objetivos principais deste trabalho consistiram em avaliar a toxicidade das ZnO-NPs com diferentes tamanhos (30 e 80-100 nm) no cladócero Daphnia magna e comparar estes efeitos com os homólogos de tamanho micrómetro (ZnO > 200 nm) e a forma iónica (ZnCl2). Os efeitos foram avaliados nos parâmetros de imobilização, inibição alimentar e reprodução.

Os resultados mostraram uma relação dose-resposta entre o decréscimo dos parâmetros avaliados e a concentração das ZnO-NPs, ZnO de tamanho micrómetro e ZnCl2 testadas. De acordo com os resultados obtidos foi possível concluir que o ZnCl2 induziu maior toxicidade aguda para a D. magna. Contudo, para a reprodução e inibição alimentar, as nanoparticulas de ZnO mostraram ter um efeito mais tóxico. Foi observado igualmente que o tamanho das nanopartículas não influenciou a toxicidade do ZnO.

Este estudo realça a importância de se estudarem os efeitos de nanoparticulas de diferentes tamanhos uma vez que este é um parâmetro-chave que deve ser considerado quando se pretende estudar a toxicidade de nanoparticulas para o ambiente.

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keywords Nanotechnology, zinc oxide nanoparticles, Daphnia magna

abstract The rapid development of nanotechnology with the consequent

increase in the production of nanoparticles and nanoproducts presents many opportunities but also many challenges. Nanotechnology has been described as a multidisciplinary field that develops a variety of engineered nanoparticles (ENPs) for medical and industrial applications.

The properties that bring to ENPs special attention for commercial products – small size, large surface area and consequently high degree of reactivity – can also make them potentially harmful for human and ecosystem health. Therefore, assessing the potential risks associated with exposure of ENPs should be considered a major research priority before they are applied in commercial products and released to the environment.

Aquatic (freshwater and marine) environment act as potential destinations for the ENPs released to the environment through direct and/or indirect sources, thus exposing aquatic organisms to high levels of pollutants.

Zinc oxide nanoparticles (ZnO-NPs) are one of the ENPs most applied in a wide range of commercial products (e.g., sunscreens, cosmetics and paints) and its production is estimated to continue to rise in the upcoming years. As a consequence, the risk of aquatic environment contamination by these ENPs will increase.

Toxicological studies have already demonstrated that nanoscale ZnO exert toxic effects in several organisms, such as crustaceans, algae and bacteria. The toxic effects of ZnO-NPs can be complex and may be dependent of several factors such as organism tested, abiotic factors (pH, salinity, water hardness, presence of natural organic matter), physico-chemical properties of NPs, adsorption phenomena, presence of other pollutants in the same environment, among others.

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The aims of this work consisted in assessing the toxicity of ZnO-NPs with two different particle sizes (30 and 80-100 nm) in the cladoceran Daphnia magna, and compare the results with the respective bulk (ZnO > 200 nm) and ionic (ZnCl2) counterparts. The effects were assessed based in a set of endpoints: immobilisation, feeding inhibition and reproduction. The results showed a dose-response relationship between all the endpoints assessed and the exposure concentrations of ZnO-NPs, ZnO micro-sized and ZnCl2 tested. According with the results obtained, it was possible to conclude that ZnCl2 induced higher acute toxicity to D. magna. However for the feeding inhibition and reproduction endpoints, ZnO nanoparticles showed to exert higher toxicity. In addition it was observed that size did not influence ZnO toxicity. This study highlights the importance to study the effects of nanoparticles with different particle sizes since this is an important parameter to be considered when analysing the toxicity of nanoparticles to the environment.

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Index

List of figures and tables ………. v

List of figures ……… v

List of tables ………. vii

1. General introduction ……… 3

1.1. Nanotechnology and Nanoparticles ……….. 3

1.2. Characteristics of nanoparticles ……… 5

1.3. Types of nanoparticles in the environment ……….. 6

1.4. Nanoparticles-environment interactions ……….. 7

1.5. Behaviour of nanoparticles in the aquatic environment ………. 8

1.6. Aims and thesis structure ………... 9

1.7. References ………... 10

2. Effect of zinc oxide nanoparticles in aquatic organisms ……… 15

2.1. Abstract ………. 15

2.2. Zinc oxide nanoparticles ……… 16

2.2.1. Applications ……….. 16

2.2.2. Synthesis of ZnO nanoparticles ……… 16

2.2.3. Release of ZnO nanoparticles to the environment ……….... 17

2.2.4. Routes of uptake and bioaccumulation of ZnO nanoparticles ………. 19

2.2.5. Effects of ZnO nanoparticles to aquatic organisms ………... 22

2.2.5.1. Toxicity to algae ……….. 23

2.2.5.2. Toxicity to aquatic organisms ………... 25

2.2.5.3. Toxicity to fish ………. 26

2.2.5.4. Toxicity to other organisms ………... 27

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3. Effect of zinc oxide nanoparticles in Daphnia magna: size dependent effects

and counterparts ………... 40

3.1. Abstract ………. 40

3.2. Introduction ………... 41

3.3. Material and methods ………. 42

Chemicals ……… 42

Preparations of suspensions ……… 43

Nanoparticles characterization ………. 43

Test organism and culture maintenance………. 43

Acute toxicity tests ………. 43

Feeding inhibition tests ……….. 44

Chronic toxicity tests ……….. 45

Statistical analysis ……….. 45

3.4. Results ……….. 46

Particle characterization of ZnO nanoparticles and respective bulk material………. 46

Acute toxicity of different sized ZnO-NPs, ZnO micro-sized and ZnCl2 to Daphnia magna ……….. 47

Exposure and post-exposure of feeding inhibition tests ……… 47

Chronic toxicity of different sized NPs, bulk counterparts and ZnCl2 to Daphnia magna ……….. 48

3.5. Discussion ………. 52

3.6. Conclusion ……… 56

3.7. References ……… 57

4. General Discussion and Conclusion ………. 63

4.1. General Discussion and Conclusion ………. 63

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List of figures and tables

List of figures

Chapter 2

Fig.1.1. Routes of ENPs aquatic environmental exposure and possible interactions with aquatic organisms after their release, from Baun et al.,

2008……….. 17

Chapter 3

Fig. 2.1. Transmission electron microscope images of zinc oxide nanoparticles of 30nm (left), 80-100nm (center) and >200nm (right) in distilled water …………. 46

Fig. 2.2. Feeding rates of D. magna during 24h of exposure and 4h of post-exposure at concentrations of Zn, ZnO-NPs and ZnO micro-sized. Black bars denote 24h of exposure and grey bars 4h of post-exposure. Data is expressed as mean values ± standard error. (*) Statistical differences at p<0.05 ... 47

Fig. 2.3. Effects of Zn, ZnO-NPs (30 and 80-100nm) and ZnO micro-sized in the number of neonates produced by D. magna. Data is expressed as mean values ± standard error. (*) Statistical differences at p<0.05 ... 51

Fig. 2.4. Body length of 21d old Daphnia magna of Zn, ZnO-NPs (30 and 80-100 nm) and ZnO micro-sized. Data is expressed as mean values ± standard error. (*) Statistical differences at p<0.05 ... 52

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List of tables

Chapter 2

Table 1.1. Predicted environmental concentrations (PECs) shown as mode (most frequent value) and as a range of the lower and upper quantiles, Q0.15 and Q0.85, for Europe and the US for different environmental compartments (base year, 2008 modified from Gottschalk et al., 2009). a represents the RQ of ZnO-NPs for STP effluents ... 18

Table 1.2. Overview of the toxicity of ZnO nanoparticles to aquatic organisms …. 30

Chapter 3

Table 2.1. Summary of the effects of all test compounds on immobilization, feeding activities and reproduction of Daphnia magna. Results are expressed as mean ± standard error; R2 is the coefficient of determination; NOEC is defined as No-observed effect concentration; LOEC is defined as Lowest observed effect concentration ... 50

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Chapter 1

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General Introduction

1. General Introduction

1.1. Nanotechnology and nanoparticles

Nanoscience is a constant growing research area with a promising future and became a field of scientific interest around the world (Bystrzejewska-Piotrowska et al., 2009; Krysanov et al., 2010).

Nanoparticles (NPs) are known as particles with at least one of their dimensions falling into the nanoscale (1-100nm) (Handy et al., 2008; Klaine et al., 2008; De Berardis et al., 2010; Fabrega et al., (in press)), or has a specific surface area by volume greater than 60 m2/cm3 according to the European Commission Recommendation of 18 October 2011 on the definition of nanomaterial (http://eurlex.europa.eu/). They can exist in different forms, spherical, tubular, or with irregular shapes (Nowack and Bucheli, 2007).

Nanoscale materials are an intermediate state between bulk and molecular materials (Moore, 2006). The ability to synthesize and manipulate nanoscale materials in this size range has being called of nanotechnology (Nowack and Bucheli, 2007). Nanotechnology presents potential opportunities that enable the synthesis of better materials and products (EPA, 2007).

Since 1990, it has been observed an exponential increase in the development of nanoproducts (products containing nanoparticles) in several industry areas. Some of these applications can bring benefits for human life style because they can be applied for medical purposes of diagnosis, imaging and drug delivery (Nel et al., 2006; Bystrzejewska-Piotrowska et al., 2009).

There are currently many types of nanoproducts applications in the market: electronics, optics, textiles, medical devices, pharmaceutics, telecommunications, cosmetics, food packaging, fuel cells, environmental remediation processes and for catalytic applications (Moore, 2006; Nowack and Bucheli, 2007; Handy et al., 2008).

With the accelerating rate on the production and use of new manufactured nanoparticles (MNPs), the release of NPs into the environment (aquatic, terrestrial and atmospheric) will occur sooner or later (Nowack and Bucheli, 2007). The production estimation of MNPs for 2004 was of 2000 tons and it is expected to increase to 58.000 tons between 2011 and 2020 according to Nowack and Bucheli (2007).

Up to now there is insufficient information about the fate, behaviour and toxicity of manufactured NPs when they reach the environment and if they pose a serious environmental and health threat (Crane et al., 2008; Krysanov et al., 2010) due to contradictory results found in the literature regarding these aspects (Dybowska et al.,

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2011). So it becomes necessary, at first, to access the risks of MNPs before they are applied in commercial products to ensure a safe manufacturing and a sustainable nanotechnology industry (Colvin, 2003).

One of the major concerns that has been raised as a consequence of this fast development relates to what is being done to access environmental risks associated with this massive production of NPs for numerous applications and their consequent release of into the environment (Dybowska et al., 2011), as well if this exponential growth of consumer products containing NPs outweighs their many benefits for the society (Colvin, 2003; Barrena et al., 2009).

It has been reported that the development of new technologies implying the use of nanoparticles has been growing much faster than the development of studies that access the implications of these materials to the environment (Krysanov et al., 2010). It should be extremely important to develop at first studies that assess the potential effects of nanoparticles especially to the aquatic environment because natural water bodies are one of the final destinations for nanoparticles through run-off, domestic/industrial wastewaters and also direct release (Baun et al., 2008).

Another concern associated with NPs applications is related to the fact that these days more and more NPs are being used as tools for nanoremediation. Nanoremediation is defined as being the use of inorganic NPs in already polluted environments aiming to reduce and/or eliminate these pollutants (Sanchez et al., 2011). This process involves some advantages (e.g., reduces the costs of clean-up of large scales and eliminates the necessity for other treatments).

However it is also necessary to take into account the costs/benefits of the application of NPs as remediation agents because of the associated risks that are inherent when NPs are applied into the environment (Sanchez et al., 2011). Nanosized materials may not migrate efficiently to be valuable for remediation (Barrena et al., 2009) thus becoming potential health hazards (Johnston et al., 2010). However, many imperative questions remain unanswered because this research area has not yet been completely examined in any great detail.

One main lack of information is coming from the fact that most studies do not present long-term effects associated with the use of NPs for environmentally polluted sites (EPA, 2007; Sanchez et al., 2011).

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1.2. Characteristics of nanoparticles

At the nano-scale level, the properties of materials differ substantially from the properties of the respective bulk material of the same composition (Xiong et al., 2011) which from the industrial and medical point of view result in performances with exceptional achievements (Nel et al., 2006). Therefore the use of NPs has obvious advantages (Tomilina et al., 2011).

On the other hand, they can become potentially harmful to the environment and living organisms causing adverse effects if NPs – biologic systems interactions occur (Nel et al., 2006).

According to Heinlaan et al., (2008) the physico-chemical differences between nanoparticles and bulk materials will induce a difference of bioavailability and toxicity of these materials.

The properties that make nano-scale materials an attraction for both industrial and medical applications but also a potential danger to living organisms are:

- Small size and large surface area;

As a consequence of their small size, NPs may become more reactive, due to its large surface area (Elsaesser and Howard, 2011), allowing them to accumulate and/or penetrate in cells and/or organisms more efficiently, possibly causing higher toxicity that would not be possible with the same material in the bulk form (Brayner et al., 2010; Peng et al., 2011). Indeed, NPs of CuO showed to be more toxic than bulk CuO for crustaceans Daphnia magna and Thamnocephalus platyurus (Heinlaan et al., 2008). However this is not a straight rule because they can form aggregates with sizes comparable with their bulk counterparts which may change their toxic potential (Wong et al., 2010).

- Shape (Bystrzejewska-Piotrowska et al., 2009);

Shape is also a relevant factor for nanoparticle toxicity. For example, carbon nanotubes are known to easily pierce cell membranes thus being able to cause toxicity (Bystrzejewska-Piotrowska et al., 2009). In a case study, Pal et al., (2007) observed that different shaped (truncated triangular nanoparticles, spherical nanoparticles and rod-shaped nanoparticles) silver nanoparticles induced different levels of bacterial growth inhibition to the gram-negative bacterium Escherichia coli. Truncated triangular silver nanoparticles were the ones to cause higher biocidal action when compared with the other shaped nanoparticles investigated (Pal et al., 2007).

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- Chemical composition (Nel et al., 2006);

Nanoparticle toxicity can be influenced by the chemical toxicity of materials from which they are made, thus making it a call for attention when introducing nanoparticles into the environment (Bystrzejewska-Piotrowska et al., 2009).

- Solubility (Nel et al., 2006);

Solubility is considered a key factor for aquatic toxicity (Kahru and Dubourguier, 2010). For instance, the solubilisation of metal-containing NPs is considered one of the main factors for nanoparticle toxicity (Kahru and Dubourguier, 2010). Indeed, Miao et al.,(2010) described that the toxic effect of ZnO-NPs to the marine diatom Thalassiosira pseudonana could be explained by the release of zinc ions to the test medium.

- Aggregation and agglomeration processes (Nel et al., 2006).

Aggregation processes often occur in aquatic environments which may lead to a different behaviour and consequent different impact to the environment (Kahru and Dubourguier, 2010).

In order to measure toxicological endpoints, the properties of nanoparticles need to be fully understood and characterized otherwise the effects of nanoparticles can be wrongly attributed to a certain property of the nanomaterial when the responsible for the effects may come from other pollutant or derived from impurities (Handy et al., 2008; Elsaesser and Howard, 2011).

1.3. Types of nanoparticles in the environment

NPs can be found in the environment resulting from natural and/or anthropogenic sources (Klaine et al., 2008). Non-natural nanoparticles can be produced unintentionally during combustion processes or intentionally produced being designated as engineered NPs (ENPs) (Nowack and Bucheli, 2007).

Natural NPs have always existed among us (Handy et al., 2008) and have been used for millions of years by humankind (Nowack and Bucheli, 2007). Natural NPs have their origin in natural sources such as volcanic ashes, byproducts of combustion fuels (e.g. coal, petroleum and wood burning) erosion processes and others (Handy et al., 2008; Bystrzejewska-Piotrowska et al., 2009). Examples of natural NPs are: organic colloids, such as humic and fulvic acids, and aerosols such as organic acids and sea salts (Nowack and Bucheli, 2007).

Contrary to natural NPs, ENPs show higher levels of complexity with unique physico-chemical features to achieve the idealized properties for the product application.

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Examples of ENPs are: fullerenes such as C60, carbon nanotubes (CNTs) and metal oxides, such as TiO2 and ZnO (Nowack and Bucheli, 2007).

There are different classes of ENPs depending on their chemical properties (Klaine et al., 2008). These classes include: carbon nanoparticles (fullerenes and nanotubes), metal oxides (ZnO, TiO2, Ce2O3), zero valent metals (Au, Ag), semiconductors (quantum dots) and nanopolymers (dendrimes) (Klaine et al., 2008). Metal oxide NPs and CNTs are the most produced classes of nanoparticles for commercial and industrial nanoproducts (Klaine et al., 2008; Johnston et al., 2010).

1.4. Nanoparticles-environment interactions

The behaviour of NPs in the environment can be complex and dependent on many processes, like abiotic factors (e.g. pH, salinity, water hardness, presence of natural organic matter), physico-chemical properties of NPs, adsorption phenomena, presence of other pollutants in the same environment, that in return may influence their toxicity (Handy et al., 2008).

Also in the environment, NPs will be in contact with different substances such as small structures (e.g., atoms, single molecules and/or macromolecules), organic natural matter, soil compounds, microbes and others that may enhance the formation of coatings which in turn may modify NPs surfaces and affect their reactivity (Handy et al., 2008).

Moreover, NPs can be released to the environment as free nanoparticles, functionalized nanoparticles, in aggregates or embedded in a matrix (Nowack and Bucheli, 2007; Bystrzejewska-Piotrowska et al., 2009). Once in the environment they can disperse into water, soil or air and act as potential environment hazards by biomagnification in the food chain (Nowack and Bucheli, 2007) affecting many groups of organisms.

Within the cells, due mostly to its small size and consequent large surface area, NPs can display a higher number of reactive oxygen species (ROS) on the surface which is currently the best-developed paradigm of NPs toxicity (Nowack and Bucheli, 2007). This event can cause damages to lipids, carbohydrates, DNA and proteins (Nel et al., 2006; Nowack and Bucheli, 2007; Elsaesser and Howard, 2011). Besides ROS production there are other main causes of nanoparticle toxicity. According to Miller et al., (2010) the dissolution of metal ions from metal oxide nanoparticles can also display an important role

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in the toxicity of NPs. ZnO-NPs has already been reported to cause toxicity through these two pathways (Franklin et al., 2007; Xia et al., 2008).

1.5. Behaviour of nanoparticles in the aquatic environment

When it comes to aquatic organisms, there are still some uncertainties about the exposure effects of NPs because it is not yet fully understood the fate and behaviour of NPs in the water column.

For instance, marine environments compared with freshwater environments are more alkaline (higher pH) and present higher ionic strength than freshwater environments (Klaine et al., 2008). As a result, NPs are likely to suffer processes of aggregation (Klaine et al., 2008) being less available to cause toxicity.

Not many studies have addressed the influence of abiotic factors such as pH, ionic strength, NOM and others in the behaviour of NPs in natural environments (Wong et al., 2010), especially in very small nanoparticles (Bian et al., 2011). The results obtained from Bian et al., (2011) showed that different conditions of ionic strength, pH, adsorption of humic acid and particle size affect the aggregation and dissolution of ZnO-NP in aqueous solutions.

Agglomeration and aggregation processes of ENPs can result in deposition of nanoscale particles in sediments (Klaine et al., 2008). However, do to their high reactivity they may not be in this state forever (Lv et al., 2012) and they can be object of geochemical processes that may disperse NPs into the water column (Klaine et al., 2008). Moreover, Yu et al.,(2011) reported that pH values closer to the zero point of charge for ZnO-NPs (9.4 – 9.5) formed aggregates, likely due to reduced repulsive interactions between particles.

Nanoparticles can also suffer action of turbulent waters, a variety of chemicals (detergents, organic matter), be coated by proteins, interact with humic and fulvic acids that may keep them disperse (Klaine et al., 2008).

In a recent study, Lv et al., (2012) reported that elements such as phosphates (major source of contamination in water) play a role in the behaviour of metal oxide NPs in aquatic environments. Due to their high specific surface, hence, high reactivity, metal oxide NPs can interact easily with phosphates influencing their speciation (Lv et al., 2012). The results of their study showed that phosphates were able to reduce the solubility of ZnO-NPs and rapidly caused aggregation (Lv et al., 2012). These findings have significant

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environmental implications since toxicity of ZnO-NPs could be greatly reduced in presence of phosphates.

Therefore, taking into account the influence of environmental factors, the behaviour of NPs in different aquatic environments is likely to be different having in return a different impact for aquatic organisms (Klaine et al., 2008).

1.6. Aims and thesis structure

The present work aimed to assess the biological effects of ZnO engineered nanoparticles with different particle sizes to the aquatic invertebrate Daphnia magna, since it is reported that the smaller the size, the higher the toxicity of NPs since they are able to interact more easily with organisms. However, as it was reported before, NPs can suffer aggregation processes once they reach the environment. Therefore, to compare the effects of two ZnO-NPs with different particle sizes (30 nm and 80-100 nm) to D. magna, we also exposed daphnids to ZnO micro-sized (>200 nm) and ZnCl2. To achieve this purpose, acute and chronic exposures were performed.

In addition, to not exclusively rely to a single aquatic species, it was elaborated a brief overview of the impacts of ZnO-NPs to aquatic organisms.

Therefore the present work was organized as follows:

- A first chapter where concepts related to nanoparticles are presented such as information about their unique features, wide range of applications and behaviour in the environment;

- A second chapter entitled “Effect of zinc oxide nanoparticles in aquatic organisms” where the effects of ZnO-NPs to some aquatic species are reported based in the literature available.

- A third chapter entitled of “Effect of zinc oxide nanoparticles in Daphnia magna: size dependent effects and counterparts”. On this chapter the effects of ZnO-NPs were assessed in a broad of several endpoints: immobilisation, feeding inhibition and reproduction.

- The fourth and last chapter provides a general discussion and conclusions of this work.

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1.7. References

Barrena, R., Casals, E., Colón, J., Font, X., Sánchez, A., Puntes, V., 2009. Evaluation of the ecotoxicity of model nanoparticles. Chemosphere 75, 850-857.

Baun, A., Hartmann, N., Grieger, K., Kusk, K., 2008. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: a brief review and recommendations for future toxicity testing. Ecotoxicology 17, 387-395.

Bian, S.-W., Mudunkotuwa, I.A., Rupasinghe, T., Grassian, V.H., 2011. Aggregation and Dissolution of 4 nm ZnO Nanoparticles in Aqueous Environments: Influence of pH, Ionic Strength, Size, and Adsorption of Humic Acid. Langmuir 27, 6059-6068.

ra ner . ahoumane . . pr mian . jediat . e er .l. out . i vet, F., 2010. ZnO nanoparticles: synthesis, characterization, and ecotoxicological studies. Langmuir 26, 6522-6528.

Bystrzejewska-Piotrowska, G., Golimowski, J., Urban, P.L., 2009. Nanoparticles: Their potential toxicity, waste and environmental management. Waste Manage. 29, 2587-2595. Colvin, V.L., 2003. The potential environmental impact of engineered nanomaterials. Nat Biotech 21, 1166-1170.

Crane, M., Handy, R., Garrod, J., Owen, R., 2008. Ecotoxicity test methods and environmental hazard assessment for engineered nanoparticles. Ecotoxicology 17, 421-437.

De Berardis, B., Civitelli, G., Condello, M., Lista, P., Pozzi, R., Arancia, G., Meschini, S., 2010. Exposure to ZnO nanoparticles induces oxidative stress and cytotoxicity in human colon carcinoma cells. Toxicol. Appl. Pharm. 246, 116-127.

Dybowska, A.D., Croteau, M.-N., Misra, S.K., Berhanu, D., Luoma, S.N., Christian, P., O'Brien, P., Valsami-Jones, E., 2011. Synthesis of isotopically modified ZnO nanoparticles and their potential as nanotoxicity tracers. Environ. Pollut. 159, 266-273.

Elsaesser, A., Howard, C.V., 2011. Toxicology of nanoparticles. Adv. Drug Deliver. Rev. EPA, 2007. Nanotechnology white paper. (100/B-07/001). United States Environmental Protection Agency.

Fabrega, J., Luoma, S.N., Tyler, C.R., Galloway, T.S., Lead, J.R., (in press). Silver nanoparticles: behaviour and effects in the aquatic environment, Environ Int (2010), doi: 10.1016/j.envint.2010.10.012.

Franklin, N.M., Rogers, N.J., Apte, S.C., Batley, G.E., Gadd, G.E., Casey, P.S., 2007. Comparative toxicity of nanoparticulate ZnO, bulk ZnO, and ZnCl2 to a freshwater

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Handy, R., Owen, R., Valsami-Jones, E., 2008. The ecotoxicology of nanoparticles and nanomaterials: current status, knowledge gaps, challenges, and future needs. Ecotoxicology 17, 315-325.

Heinlaan, M., Ivask, A., Blinova, I., Dubourguier, H.-C., Kahru, A., 2008. Toxicity of nanosized and bulk ZnO, CuO and TiO2 to bacteria Vibrio fischeri and crustaceans Daphnia magna and Thamnocephalus platyurus. Chemosphere 71, 1308-1316.

Johnston, B.D., Scown, T.M., Moger, J., Cumberland, S.A., Baalousha, M., Linge, K., van Aerle, R., Jarvis, K., Lead, J.R., Tyler, C.R., 2010. Bioavailability of nanoscale metal oxides TiO2, CeO2, and ZnO to fish. Environ. Sci. Technol. 44, 1144-1151.

Kahru, A., Dubourguier, H.-C., 2010. From ecotoxicology to nanoecotoxicology. Toxicology 269, 105-119.

Klaine, S.J., Alvarez, P.J.J., Batley, G.E., Fernandes, T.F., Handy, R.D., Lyon, D.Y., Mahendra, S., McLaughlin, M.J., Lead, J.R., 2008. Nanomaterials in the environment: behavior, fate, bioavailability, and effects. Environ. Toxicol. Chem. 27, 1825-1851.

Krysanov, E., Pavlov, D., Demidova, T., Dgebuadze, Y., 2010. Effect of nanoparticles on aquatic organisms. Biology Bulletin 37, 406-412.

Lv, J., Zhang, S., Luo, L., Han, W., Zhang, J., Yang, K., Christie, P., 2012. Dissolution and Microstructural Transformation of ZnO Nanoparticles under the Influence of Phosphate. Environ. Sci. Technol. 46, 7215-7221.

Miao, A.-J., Zhang, X.-Y., Luo, Z., Chen, C.-S., Chin, W.-C., Santschi, P.H., Quigg, A., 2010. Zinc oxide engineered nanoparticles dissolution and toxicity to marine phytoplankton. Environ. Toxicol. Chem. 29, 2814-2822.

Miller, R.J., Lenihan, H.S., Muller, E.B., Tseng, N., Hanna, S.K., Keller, A.A., 2010. Impacts of metal oxide nanoparticles on marine phytoplankton. Environ. Sci. Technol. 44, 7329-7334.

Moore, M.N., 2006. Do nanoparticles present ecotoxicological risks for the health of the aquatic environment? Environ. Int. 32, 967-976.

Nel, A., Xia, T., Madler, L., Li, N., 2006. Toxic potential of materials at the nanolevel. Science 311, 622-627.

Nowack, B., Bucheli, T.D., 2007. Occurrence, behavior and effects of nanoparticles in the environment. Environ. Pollut. 150, 5-22.

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Pal, S., Tak, Y.K., Song, J.M., 2007. Does the Antibacterial Activity of Silver Nanoparticles Depend on the Shape of the Nanoparticle? A Study of the Gram-Negative Bacterium Escherichia coli. Appl. Environ. Microb. 73, 1712-1720.

Peng, X., Palma, S., Fisher, N.S., Wong, S.S., 2011. Effect of morphology of ZnO nanostructures on their toxicity to marine algae. Aquat. Toxicol. 102, 186-196.

Sanchez, A., Recillas, S., Font, X., Casals, E., Gonzalez, E., Puntes, V., 2011. Ecotoxicity of, and remediation with, engineered inorganic nanoparticles in the environment. Trac-Trends in Analytical Chemistry 30, 507-516.

Tomilina, II, Gremyachikh, V.A., Myl'nikov, A.P., Komov, V.T., 2011. Changes in biological characteristics of freshwater heterotrophic flagellates and cladocerans under the effect of metal oxide nano- and microparticles. Inland Water Biology 4, 475-483.

Wong, S.W.Y., Leung, P.T.Y., Djurisic, A.B., Leung, K.M.Y., 2010. Toxicities of nano zinc oxide to five marine organisms: influences of aggregate size and ion solubility. Analytical and Bioanalytical Chemistry 396, 609-618.

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Xiong, D., Fang, T., Yu, L., Sima, X., Zhu, W., 2011. Effects of nano-scale TiO2, ZnO and their bulk counterparts on zebrafish: acute toxicity, oxidative stress and oxidative damage. Sci. Total. Environ. 409, 1444-1452.

Yu, L.-p., Fang, T., Xiong, D.-w., Zhu, W.-t., Sima, X.-f., 2011. Comparative toxicity of nano-ZnO and bulk ZnO suspensions to zebrafish and the effects of sedimentation, (center dot) OH production and particle dissolution in distilled water. Journal of Environmental Monitoring 13, 1975-1982.

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Chapter 2

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Effect of zinc oxide nanoparticles in aquatic organisms

2. Effect of zinc oxide nanoparticles in aquatic organisms

2.1. Abstract

The increasing use of engineered nanoparticles (ENPs) in industrial and daily life applications will result in the release of these nanoscale materials into environmental compartments.

The aquatic environment is one of the possible and last sinks for any chemical/pollutant that reaches the environment. For this reason, aquatic organisms, especially invertebrates, are widely used in toxicity testing.

It is demonstrated that the behaviour of nanoparticles in the environment is still rather poorly understood, requiring closer attention from regulatory agencies since the release of ENPs into environmental compartments will increase in the very near future.

One of the most produced classes of NPs is the class of metal oxide nanoparticles such as ZnO nanoparticles (ZnO-NPs) due to their application diversity. As a consequence of their wide use they will inevitably reach aquatic ecosystems.

Measurements of environmental concentrations of nanoparticles in the environment are difficult to quantify. However, sophisticated probabilistic methods showed predicted environmental concentrations (PECs) for nano-ZnO in U.S and Europe in the range of µg.L-1 for waters to mg.Kg-1 for soils in different environmental compartments.

Knowing that ZnO-NPs pose a risk for aquatic organisms, the aim of this study was to summarize the present knowledge on the fate, behaviour, uptake routes and biological effects of ZnO-NPs to aquatic organisms.

Lastly, current knowledge gaps are pointed out and brief recommendations for future developments are made.

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2.2. Zinc oxide nanoparticles

2.2.1. Applications

ZnO-NPs are considered a versatile and technologically important material (Meulenkamp, 1998) because it can be found in a wide range of applications such as biosensors, electronic materials (Brayner et al., 2010), ceramics, rubber manufacturing, as a fungicide (Naddafi et al., 2011), in wastewater treatments (Wong et al., 2010), coatings, paints (Blinova et al., 2010) and textile industry (Heinlaan et al., 2008).

One of the unique properties of ZnO-NPs relates to its large UV spectrum of attenuation properties (Handy et al., 2008) making them one of the most used nanoparticles in personal care products (e.g., sunscreens, tooth pastes, cosmetics) (Blinova et al., 2010; Xiong et al., 2011). This brings in return great attention because of its high possibility to be released into the environment (Franklin et al., 2007; Wong et al., 2010). Researchers estimated that at least 25% of the total amount of sunscreen applied in the skin is washed away during bathing recreations (Wong et al., 2010).

Since ZnO-NPs are used in a wide range of commercial products, this occurrence leads to an increased interest in the behaviour and possible toxic impacts of such materials in the aquatic environment.

2.2.2. Synthesis of ZnO nanoparticles

There are several physical and chemical methods used for the manufacturing of ZnO-NPs. ZnO-NPs can be synthesized by various approaches including sol-gel processing, homogeneous precipitation, mechanical milling, organometallic synthesis, microwave method, spray pyrolysis, thermal evaporation, mechanochemical synthesis (Hong et al., 2009) and forced hydrolysis in DEG, di(ethylene glycol) medium (Brar et al., 2010).

However it has to be taken into account that NPs tend to aggregate due to its large surface area (as a consequence of their small size) (Hong et al., 2009). So, in order to improve the dispersion of NPs, it is necessary to modify their surfaces (Hong et al., 2009). Examples of dispersants that help to maintain the stability of suspensions of ZnO-NPs are Polyethylene Glycol (PEG), Polyvinylpyrolidone (PVP) (Zhang et al., 2007), tri-n-octylphosphine (TOPO), sodium dodecyl sulfate (SDS) and others (Brayner et al., 2006).

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2.2.3. Release of ZnO nanoparticles to the environment

The increasing use of nanoparticles will lead to the intentionally/unintentionally introduction of NPs in the environment. NPs can remain in the environment for long periods of time and become potentially toxic to aquatic environments because once they are released they may interact with aquatic surfaces, biological species or experience aggregation and sedimentation processes (Brar et al., 2010).

Due to the diversity of ZnO-NPs and other ENPs applications, they can enter the environment through many pathways by both point and non-point sources (Wiechers and Musee, 2010). The main release sources of ENPs into the environment include accidental spillages from industrial activities (e.g., release of liquid and solid waste streams) and

Figure 1.1. Routes of ENPs aquatic environmental exposure and possible interactions with aquatic organisms after their release, from Baun et al., (2008a).

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during transportation (Wiechers and Musee, 2010). Un-removed NPs from wastewater treatments and agricultural activities (e.g., use of municipal sludge, pesticides) are also sources of release of ENPs into the environment (Wiechers and Musee, 2010).

Figure 1.1 shows the possible routes of ENPs environmental exposure after their release into the aquatic environment. Once they are released ENPs can be suspended in the water column and be taken by planktonic (e.g., daphnids), or by sediment dwelling invertebrates since the sediments are also considered as a potential sink for many contaminants in water ecosystems thus being subject to high levels of contaminants (Baun et al., 2008a).

Given the fact that the release of ENPs from products constitutes a pathway for these nanoparticles to reach the environment, measurements of environmental concentrations of ENPs should be widely examined.

There is few data available reporting the presence of nanoparticles in the environmental compartments, however for ZnO-NPs, Gottshalk et al.,(2009) suggests that several ENPs, including ZnO-NPs may be present in different environmental compartments (Table 1.1).

Table 1.1. Predicted environmental concentrations (PECs) for ZnO-NPs shown as mode (most frequent value) and as a range of the lower and upper quantiles, Q0.15 and Q0.85, for Europe and

U.S. for different environmental compartments (base year, 2008 modified from Gottschalk et al.,(2009). a represents the RQ of ZnO-NPs for STP effluents.

Europe U.S. Mode Q0.15 Q0.85 Mode Q0.15 Q0.85 Soil 0.093 0.085 0.661 0.050 0.041 0.274 Δμg kg-1 y-1 Sludge treated soil 3.25 2.98 23.1 1.99 1.62 10.9 Δμg kg-1y-1 Surface water 0.010 0.008 0.055 0.001 0.001 0.003 μg L -1 STP effluent 0.432 (10.8) a 0.340 1.42 0.3 (7.7)a 0.22 0.74 μg L-1 STP sludge 17.1 13.6 57.0 23.2 17.4 57.7 mg kg -1 Sediment 2.90 2.65 51.7 0.51 0.49 8.36 Δμg kg-1y-1 Air <0.0005 <0.0005 μg m-3

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Gottschalk et al., (2009) aimed to calculate predicted environmental concentrations (PECs) of several NPs (TiO2, ZnO, Ag, CNT and fullerenes) based on a probabilistic material flow analysis from a life-cycle perspective of products containing ENPs. In a second aim, to assess risk quotients (RQs) posed by ENPs, the simulated PECs where compared to the predicted no effect concentration (PNEC) based on ecotoxicological data from the literature for the each environmental compartment.

Table 1.1 shows PECs of ZnO-NPs for air, surface water, sewage treatment plant (STP) effluent and sewage sludge and simulation of the amount of ZnO-NPs deposited in soil, sludge-treated soil and sediment in 2008. For soils and sediments, using estimations of the worldwide market evolution for products containing ZnO-NPs for the period 2001-2012 and assuming zero concentrations in 2000, authors were able to roughly estimate the amount of ZnO-NPs deposited in these compartments for each year of the period considered.

Of all ENPs considered in the study, ZnO-NPs showed to be one of the ENPs with highest concentrations in all compartments with modeled concentrations for natural surface waters of 0.010 µg.L-1 and 0.432 µg.L-1 for treated wastewater in Europe. These results reflect that ZnO-NPs may cause significant risks to the aquatic environment.

The study indicated that the risk for aquatic organisms when exposed to ZnO-NPs emanates mostly from sewage treatment effluents due to the fact the risk quotient (RQ) were greater than critical value of one (Table 1.1.) for this compartment but lower to the other compartments (Gottschalk et al., 2009).

Therefore it is suggested that more investigations should be performed in order to evaluate the real risks posed to aquatic organisms by ZnO-NPs.

2.2.4. Routes of uptake and bioaccumulation of ZnO nanoparticles

Uptake of nanoparticles into aquatic biota has been reported to be through direct ingestion and/or entry across epithelial surfaces (e.g., gills, olfactory organs or body wall) (Moore, 2006).

In addition to the uptake, the increase in the concentration of a chemical in a biological organism over time (bioaccumulation) has also to be considered when evaluating potential hazards and risks of NPs since they may become precursors of toxicity (Fabrega et al., (in press)). One important question concerning bioaccumulation is whether NPs bioaccumulate after they penetrate into the organism or if they only stay adsorbed to external surfaces causing cell damage.

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In a pioneering study with bacteria, Brayner et al., (2006) observed, through TEM images, either accumulation of ZnO nanoparticles in the bacterial membrane of E.coli as well as internalization of these nanoparticles as a result of cell wall disorganization. Kumar et al., (2011a) also observed uptake and consequent internalization of ZnO and TiO2 nanoparticles in Salmonella typhimurium.

According to Moore (2006), the uptake of NPs at the cellular level, is thought to be through endocytosis. In endocytosis, molecules or particles between 1 and 100nm are taken up by invagination of the plasma membrane leading to the formation of vesicles that encloses the material and transports it into the cell.

For aquatic invertebrates, Zhu et al., (2009b) studied the toxicity of six NPs (i.e., ZnO, TiO2, Al2O3, C60, SWCNTs and MWCNTs) through an acute bioassay, using immobilisation and mortality as endpoints in Daphnia magna.

One of the aims of the study was to create dose-dependency curves for all nanoparticle water suspensions and then to compare with the toxicity of respective bulk counterparts to determine if the size of NPs affected their toxicity to D. magna. In addition, the present study also aimed to evaluate the uptake of NPs by D. magna.

ZnO-NPs suspension showed to be the most toxic to D. magna among all NPs studied with an EC50 of 0.622 mg.L-1 and an LC50 of 1.511 mg.L-1.

The uptake of nanoparticles by D. magna was recorded by using a microscope with a digital camera (2 x 10 resolution). The authors reported that after 48h of exposure ingestion of NPs and consequent accumulation in the gut occur at the highest concentrations.

There is also available data for other ENPs. For instance, Heinlaan et al., (2011), studied the changes in the midgut of D. magna, using TEM technique, when exposed to CuO-NPs and CuO bulk in an attempt to evaluate the size-related effects. TEM is considered an essential tool for (nano)ecotoxicological studies because it enables the visualization of ultrastructural changes of cells and tissues when exposed to nanoparticles (Heinlaan et al., 2011).

At the end of 48h of exposure, CuO bulk particles at a nominal concentration of 175 mg.CuO.L-1 (48h-EC50) were found wrapped in aggregates by the peritrophic membrane (TPM). This membrane is designed to protect the gut epithelium and also regulate nutrient and enzyme exchanges (Heinlaan et al., 2011).

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Contrary to CuO bulk, CuO-NPs at a nominal concentration of 4 mg.CuO.L-1, corresponding to the 48h-EC50 value for CuO-NPs, nanoparticles were taken up by D. magna within 10 min of exposure and were found dispersed in the gut of D. magna along with bacterial colonization.

Authors suggested that the appearance of bacteria could be caused by suppression of the immune response of D. magna. Since no bacterial colonization took place during exposure of D. magna to CuO bulk it is suggested that the presence of bacteria observed during exposure of CuO-NPs can be a specific mode of toxicity induced by these nanoparticles.

Taking into account the studies mentioned above, their observations come to reinforce the fact that the potential impacts of ENPs for aquatic organisms should not be neglected since exposure of these nanoparticles may pose a risk of bioaccumulation which may lead to dead of organisms such as filter-feeding organisms (e.g., Daphnia magna) affecting like this the balance of the aquatic environment.

Another exposure pathway that may lead to the uptake of ENPs by aquatic organisms can be the through the sorption of NPs to algae. For instance, Croteau et al., (2011) showed the behaviour of the freshwater snail, Lymnaea stagnalis, when exposed to well-characterized isotopically modified ZnO-NPs (67ZnO nanoparticles).

The aims of the study were to address the uptake rates of 67Zn in the form of ZnO-NP when ingested with food and the uptake rates of 67Zn also when ingested with food. Benthic diatom Nitzschia palea was used as food source.

Whether the form of 67Zn, the results showed that both forms were efficiently assimilated by L. stagnalis showing that 67Zn from ZnO-NPs appear equally bioavailable as the ionic 67_Zn.

In another phase of the study, authors observed that high concentrations of ZnO-NPs caused interference in the feeding rates of L. stagnalis and also affected their defecation rates at high concentrations (106 nmol g-1) of ZnO-NPs, thus showing evidences of dietary metal stress. For the authors the reduced feeding activity observed by L. stagnalis remained unclear if it was due to high concentrations of Zn (ionic) derived from ZnO-NPs or due to the ZnO-NPs themselves.

To date, many studies have focused especially on the direct effect of NPs. However, NPs can also act as contaminant carriers for other bioavailable toxicants (e,g., metal-containing NPs, C60) (Baun et al., 2008b; Croteau et al., 2011).

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It is known that metal-containing NPs are already widely used in consumer products. Therefore, after their discharge to the environment, these NPs may interact with other pollutants (Baun et al., 2008b). This interaction may lead to a change on the bioavailability of these pollutants to aquatic organisms (Baun et al., 2008b) making it a point of concern in the future.

For instance, in a recent study, Naddafi et al., (2011) studied the bioavailability of phenanthrene adsorbed to ZnO-NPs using Daphnia magna as the organism test.

The results showed that the bioaccumulation of phenanthrene in D. magna was enhanced by the presence of ZnO-NPs when compared with phenanthrene free of ZnO-NPs since its toxicity for 24h and 48h was 1.7 and 2.1 times higher, respectively (Naddafi et al., 2011).

Therefore, and based on the previous study, when it comes to risk assessments of nanoparticles, it is then important to consider not only the inherent toxicity of NPs but also the possible interactions between NPs with already existing environmental contaminants.

Also, the increasing release of ENPs into the environment, their low biodegradability and possible association with the feeding habits of aquatic invertebrates, calls the urge for faster development studies to better understand the behaviour and mechanisms of uptake/bioaccumulation of ENPs in aquatic ecosystems because as described in this section ENPs are easily taken up by these organisms which in the worst cases may lead to death.

2.2.5. Effects of ZnO nanoparticles to aquatic organisms

The impacts of common metals for the health and the environment are well known (Brar et al., 2010). However, when metals take the form of nanoparticles, the potential hazards due to their shape and size are yet to be explored (Brar et al., 2010).

Many studies have already investigated the toxicity of ZnO-NPs on several aquatic organisms (Table 1.2), however the toxic mechanisms are still unclear (Bai et al., 2010).

In studies where the toxicity and interactions of ZnO-NPs with aquatic organisms are assessed, they are in most of the cases compared to either larger ZnO-NPs (ZnO micro-sized) or to ionic Zn (Franklin et al., 2007; Heinlaan et al., 2008; Zhu et al., 2008; Aruoja et al., 2009; Wiench et al., 2009; Zhu et al., 2009a; Zhu et al., 2009b; Wong et al., 2010; Fairbairn et al., 2011; Yu et al., 2011).For metal-based nanoparticles it is important to take into consideration the solubility of ions on their toxicity since many studies have

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considered Zn ions as one of the main key factors that accounts for the toxicity of ZnO-NPs to aquatic organisms(Franklin et al., 2007; Heinlaan et al., 2008; Aruoja et al., 2009).

ZnO nanoparticles are widely incorporated in commercial merchandise. However, their environmental impact and their mechanisms of toxicity are not yet fully understood. According to the available literature the toxicity of ZnO-NPs to aquatic organisms is in some part attributed to its solubility in water but also with ROS production even if the this last one still remains a open question (Bai et al., 2010; Miller et al., 2010).

In the next section, several studies concerning the effects of ZnO-NPs to aquatic organisms of different trophic levels of the food chain will be presented.

2.2.5.1. Toxicity to algae

For algae species the endpoint most assessed is the growth inhibition endpoint (Table 1.2.).

Franklin et al.,(2007) developed a study where it presented the effects of ZnO-NPs (nominal particle size of 30 nm), ZnO bulk and ZnCl2 in the growth rate of the freshwater algae Pseudokirchneriella subcapitata. The results showed that the toxicity of ZnO-NPs did not present statistical differences when compared with ZnO bulk and ZnCl2, showing 72h-IC50 values of 68, 63 and 61 µg Zn2+.L-1, respectively.

In addition, concentrations of dissolved Zn2+ ions from ZnO-NPs were determined by equilibrium dialysis, with a pore size of about 1 nm (permeable to Zn2+ ions but not to ZnO particles)and compared with the concentration of dissolved Zn2+ ions from bulk ZnO and ZnCl2. The results showed that at pH 7.6 rapid dissolution rate for both ZnO-NPs and ZnO bulk occurred within 6h yielding similar dissolved zinc concentrations. At the end of the 72h of the experiment, 19% of the nominal concentration (100 mg.L-1) of ZnO-NPs and ZnO bulk was dissolved. Taking into account the dialyzed zinc concentrations obtained for all zinc compounds during the experiment, the 72h-IC50 were calculated. Once again the 72h-IC50 values showed similar toxicities.

Therefore authors suggested that toxicity of ZnO particles (ZnO-NPs and bulk ZnO) as well as ZnCl2 to P. subcapitata could be essentially due to dissolved zinc.

In a similar study, Arouja et al., (2009) also investigated the toxicity of ZnO-NPs (50-70 nm particles size) to P. subcapitata and aimed to clarify if the influence of particle size and Zn2+ ions dissolution played an important role for the toxicity by using other zinc compounds (bulk ZnO and ZnSO4).

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Both ZnO-NPs and bulk ZnO showed to be toxic to P. subcapitata at low concentrations (<0.1 mg Zn2+.L-1) and caused 100% of inhibition at concentrations of 0.16 mg Zn2+.L-1. As it happened in the previous study, the toxicity of ZnO-NPs and bulk ZnO showed similar toxicity (no statistical differences regarding the particle size). However,bulk ZnO showed to cause a slightly higher inhibition effect to P. subcapitata than ZnO-NPs with 72h-IC50 values for ZnO-NPs and ZnO bulk of 0.042 and 0.037 mg Zn2+.L-1, respectively.

Regarding ZnSO4, the IC50 value (calculated on metal basis) for this compound was of 0.042 mg Zn2+.L-1. Therefore and taking into account a previous study developed in their laboratory which indicated that at already low concentrations (0.1 mg.L-1) of both nano and bulk, a high fraction (between 69% and 97%) of Zn (ionic) was already bioavailable, they attribute the toxicity of ZnO-NPs to dissolved zinc ions.

Peng et al., (2011) studied the influence of different sized and shaped (sphered and rod-shaped particles) ZnO-NPs on the growth of three marine diatoms, Thalassiosira pseudonana, Chaetoceros gracilis and Phaeodactylum tricornutum, as it has been shown that dissolution reflected as toxicity, is influenced by nanoparticles’ morphologies. The particles’ dimension analysed by TEM showed that nZnO spheres ranged from 6.3 nm to 15.7 nm and nZnO rod-shaped ranged between 242 nm to 862 nm.

The extent of dissolution in seawater, measured by GFAAS (graphite furnace atomic absorption spectrometry), showed that the solubility of Zn in sphered NPs was higher (but not statistically different) when compared with rod-shaped NPs.

These observations are probably due to the differences in NPs morphology because, according to Borm et al., (2006), smaller particles due to their surface curvature and thinner diffusion layers, will reach faster equilibrium dissolution rates.

They also observed that an increase of the concentration of ZnO-NPs did not result necessarily in an increase of the amount of Zn ions in solution at equilibrium, which allowed the authors to conclude that ZnO-NPs particle concentrations did not have much influence in the dissolution behaviour of nanoparticles probably due to the observation of aggregation of NPs during the experiment.

Regarding the toxicity of ZnO-NPs to the algae species, growth of all marine species was affected by all the concentrations and all different morphologies of ZnO-NPs tested, with P. tricornutum being the less sensitive exhibiting a slow but continuous growth rate in the presence of ZnO-NPs. The differences in the sensitivity are explained by the fact that P. tricornutum has on its morphology frustules containing less silica than those of the other marine algae species tested. Silica is an important component of diatoms cell wall that

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influences their growth. However, even if Zn inhibited diatom growth and frustule formation in P. tricornutum, this diatom specie needs less quantities of Si to fulfill its requirements for the formation of its frustule when compared with the other diatom species tested. Therefore, P. tricornutum was able to exhibit continuous growth even if in a slow rate.

2.2.5.2. Toxicity to aquatic invertebrates

According to legislative organizations (e.g. REACH), the number of vertebrate animals in toxicological tests should be reduced and alternative testing approaches should be used. Therefore, aquatic invertebrates are being the model organisms most used for (nano)ecotoxicological studies since they represent an important level in the food chain of marine and freshwater ecosystems (Baun et al., 2008a). For aquatic invertebrates the immobilization endpoint is the most common endpoint studied.

There has been a wide range of results concerning the toxicity of ZnO-NPs to aquatic invertebrates.

For instance, Heinlann et al., (2008) using a recombinant Zn-sensor bacteria to compare the toxicity of Zn2+ (in the form of ZnSO4), nano ZnO and bulk ZnO, showed that the effect of all Zn compounds for crustaceans Daphnia magna and Thamnocephalus

platyurus were similar (Table 1.2), attributing this to the concentration of soluble Zn2+.

Contrary, Zhu et al., (2009b) also studied the effect of ZnO-NPs in Daphnia magna (Table 1.2). However, the value of 48h-LC50 (1.511 mg.L-1) obtained was higher than the 48h-LC50 value (3.2 mg.L-1) for ZnO-NPs reported by Heinlaan et al.,(2008).

The differences of the results from both studies could be explained by different approaches adopted in the ecotoxicological protocols since the experimental setup of the first study was performed in the dark which could have influenced the toxicity of ZnO-NPs to the organisms.

In a more recent study, Zhao Hai-zhou., (2012) assessed the effect of ZnO-NPs on the survival, reproduction and feeding behaviour of D. magna. The results obtained were in accordance with the study reported by Zhu et al., (2009b) obtaining a 48h-LC50 value for ZnO-NPs of 1.48 mg.L-1 for the mortality endpoint. For the other endpoints,

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concentration-dependent doses were observed. Possible, the reduction of food uptake, may have in turn affected growth and reproduction of D. magna.

2.2.5.3. Toxicity to fish

Studies investigating effects in vivo of ZnO-NPs in fish are still scarse (Zhu et al., 2008; Zhu et al., 2009a; Xiong et al., 2011; Yu et al., 2011) and they have been especially focused on the toxicity effects at early developmental stages (Yu et al., 2011) (Table 1.2).

Zhu et al., (2008) studied the effects of ZnO-NPs (20 nm) to zebrafish (Danio rerio) in several endpoints such as embryo survival and hatching rate and after compared with ZnO micro-sized particles (1µm).

Embryo survival and hatching rate of zebrafish showed a dose-dependency with the increase of ZnO-NPs concentrations. However, no statistical differences were found between Zn compounds for both endpoints, recording a 96h-LC50 value of 1.793 mg.L-1 for ZnO-NPs and 1.550 mg.L-1 for ZnO bulk for embryo survival and a 84h-EC50 value of 2.065 mg.L-1 for ZnO-NPs and 2.066 mg.L.-1 for ZnO bulk for hatching rate.

Dissolution of Zn2+ ions from ZnO-NPs and ZnO bulk were assessed so that toxicological tests were performed again for the same endpoints; results showed that Zn2+ concentrations affected the survival rates in 86.7% at the end of 96h and 90% of the hatching rates at the end of 84h.

It was then clear that the release of zinc ions contributed for the toxic effects in zebrafish development. However, authors reported that toxicity of Zn2+ was significantly lower that the toxicity of ZnO-NPs, therefore suggesting that maybe there are others factors responsible for the toxic effects, such as ROS production.

The ecotoxicological tests developed by these authors were performed in Mili-Q water. This fact makes us raise the question if zebrafish embros could survive during 96h only in Mili-Q water.

Later, reported by the same author, Zhu et al.,(2009a) assessed the influence of micro-sized ZnO-NPs during 96h to the embryonic development of zebrafish (Danio rerio), reporting a 84h-EC50 value (for embryo hatching) of 23.06 mg.L-1 (Table 1.2).

In this study they also measured the release of Zn2+ ion concentrations of all test solutions, by GFAA, to determine whether zinc played an important role to the development of D. rerio. By exposing zebrafish embryos to the concentrations of Zn2+

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released from ZnO-NPs, it was observed that zinc ions did not cause apparent toxicity to zebrafish when compared with the same concentrations of ZnO-NPs, which again implied that Zn2+ could not have been the only toxic agent to zebrafish.

Yu et al., (2011) (Table 1.2.) studied the effects of ZnO-NPs (30 nm) in adult zebrafish and compared the effects of ZnO-NPs with their bulk counterpart (500 nm) and Zn2+ ions, in the form of ZnSO4. The 96h-LC50 value obtained for ZnO-NPs was 3.70 mg.L -1_{. ZnO bulk suspensions showed to be more toxic than ZnO nanoparticles, however not} statistically different, with a 96h-LC50 value of 2.53 mg.L-1, probably due to aggregation of ZnO-NPs suspensions as reported by authors. For Zn2+, the toxicity to zebrafish showed to be higher when compared with the ZnO particles (96h-LC50 = 7.480 mg.L-1).

Therefore authors suggested that dissolved zinc from ZnO suspensions may have contributed for the acute toxicity to zebrafish and that the mechanisms of toxicity of Zn2+ may be different from ZnO particles. OH- generation was also determined but it was concluded they played a small role in the toxicity to zebrafish since ZnO-NPs and ZnO bulk presented very different abilities to generate OH- thus not being the main factor for toxicity.

Xiong et al., (2011) recently reported dose-dependency toxicity for ZnO-NPs (30 nm), ZnO bulk (500 nm) and Zn2+ (in the form of ZnSO4) in adult zebrafish. 96h-LC50 values for all Zn compounds obtained were 4.26 mg.L-1 to ZnO-NPs and 3.31 mg.L-1 to ZnO bulk and 8.06 mg.L-1 to Zn2+ ions, with statistical differences between ZnO-NPs and Zn2+ and between ZnO bulk and Zn2+ (Table 1.2).

Like the previous study and given the results, it was suggested that the dissolution of zinc ions it may contribute for the toxicity of ZnO-NPs and ZnO bulk but it was not considered as the main lethal mechanism for ZnO suspensions.

2.2.5.4. Toxicity to other organisms

There are also other studies concerning the effects of ZnO-NPs to organisms such as bacteria, amphibians and plants but at a lower scale.

Due to its antimicrobial activity, studies on the effects of ZnO-NPs to different bacteria species have already been documented.